Alginate-like polymers from full-scale aerobic granular sludge: content, recovery, characterization, and application for cadmium adsorption

Aerobic granular sludge (AGS) is a proven resource for the recovery of biopolymers like alginate-like polymers (ALP). This is the first report on the dynamics of ALP produced by AGS (ALP-AGS) in a full-scale wastewater treatment plant (WWTP), optimization of ALP recovery from AGS, and adsorption of cadmium (Cd2+) by ALP. Recovery of ALP was highest when using 120 mL of 0.2 M Na2CO3 at 70 °C for 45 min. Seasonal (1.5 years, over 3100 cycles) and intra-cycle changes in ALP-AGS in the WWTP were monitored. The ALP content in AGS increased in the transition period between winter and spring, reaching over 150 mg/g MLSS. In the batch reactor cycle, the ALP-AGS level peaked 2 h after the start of aeration (mean peak level: 120 mg/g MLSS), then decreased about two-fold by the end of the cycle. The ALP-AGS had a small surface area and a lamellar structure with crystalline outgrowths. The optimal conditions of Cd2+ adsorption with ALP were a dosage of 7.9 g d.m./L, a pH of 4–8, and an equilibrium time of 60 min. Carboxyl and hydroxyl groups were the key functional groups involved in Cd2+ adsorption. According to the Sips model, the maximum Cd2+ adsorption capacity of ALP-AGS was 29.5 mg/g d.m., which is similar to that of commercial alginate. AGS is a richer source of ALP than activated sludge, which ensures the cost-effectiveness of ALP recovery and increases the sustainability of wastewater treatment. Information on the chemical properties and yields of ALP from full-scale WWTPs is important for downstream applications with the recovered ALP.

www.nature.com/scientificreports/ and the waste sludge from the secondary clarifiers were collected in a sludge thickener and then dewatered using a mechanical belt thickener. The cycle of a biological reactor begins with simultaneous, piston feeding of raw wastewater and withdrawal of effluent. The whole reactor cycle lasted for 4.8 h and consisted of 0.8 h of anoxic filling with simultaneous wastewater withdrawal, an aeration period of 3.4 h, and 0.6 h of settling and biomass discharge. The volumetric exchange rate was about 25%. The AGS was sampled from February 2018 to August 2019 from both GSBRs at the end of the aeration phase in the cycle (every 25 days on average). Intra-cycle measurements of ALP content in AGS were performed in two independently operated full-scale GSBRs during the aeration phase of cycle 916th. Before analysis, the sludge was stored at 4 °C for no longer than 24 h. In the investigated period, the average biomass concentrations were 6.8 g MLSS/L (GSBR#1) and 7.0 g MLSS/L (GSBR#2). The sludge volumetric index in both GSBRs was at a level of 50 mL/g MLSS.
ALP extraction from AGS. The content ALP in AGS was measured according to Lin et al. 17 , but the procedure was optimized to ensure the highest yield of ALP from AGS. Therefore, 2.5 g of MLSS was homogenized using Ultra Turrax T25 basic (IKA-WERKE) for 5 min at 9.500 1/min, and then incubated from 15 to 60 min at temperatures varying from 50 to 80 ºC in 0.2 M Na 2 CO 3 -a volume of solution from 80 to 400 mL. Biomass was then centrifuged at 12,000 rpm for 20 min. The reaction of the supernatant was adjusted to pH 2 using 0.1 M HCl. The supernatant was centrifuged at 12,000 rpm for 20 min and the ALP was dissolved in 0.1 M NaOH. ALP was precipitated using cold EtOH adjusting the solution to 80% (v/v), lyophilized, and weighed. The extracted ALP was used in further experiments. Adsorbent preparation. Cd 2+ adsorption experiments were performed with ALP-AGS which was in the form of beads. Sodium alginate (ALG) purchased from Sigma-Aldrich was used as a reference adsorbent. The beads were prepared by dropwise addition of a viscous 2.5% (w/w) ALG or ALP-AGS aqueous solution to 0.03 M CaCl 2 solution with the help of an injection syringe. The gelation process was continued for 30 min. During the gelation process, the alginate reacts with the Ca 2+ ions from the CaCl 2 solution to form a cross-linked Caalginate. The obtained beads were left in the solution for 24 h, and then kept in distilled water at 4 °C in the refrigerator. The experiments on Cd 2+ adsorption were performed with fresh beads.
Batch adsorption experiments. Commercial stock standard Cd 2+ solution at a concentration of 1000 mg Cd 2+ /L (Sigma-Aldrich) was used. Working standards of required concentration were prepared by diluting the stock solution in distilled water. STEP 2. EvaluaƟon of ALP-AGS dynamics in a real wastewater treatment system Purpose: to define the best periods for recovery of ALP from AGS two biological rectangular reactors with aerobic granules (GSBRs) seasonal and intra-cycle measurements of ALP content in AGS STEP www.nature.com/scientificreports/ Cd 2+ removal from aqueous solution was conducted at room temperature, in a function of adsorbent dosage (0.7, 2.6, 5.3, 7.9, 10.6, 13.2, and 15.9 g d.m./L), contact time (5, 15, 30, 60, 90, 120, 150, 180 min.), pH (2,3,4,5,6,8) and initial Cd 2+ concentration (5,10,15,30,50,75,100,150,200 mg/L). The experiments were conducted in 125-mL Erlenmeyer flasks. The adsorbent dosage for pH and contact time experiments was set to 2.6 g d.m./L, while for the effect of initial Cd 2+ concentration it was fixed to 7.9 g d.m./L. The pH of the Cd 2+ working solution in the experiments, except for the effect of pH, was adjusted to 5.0 ± 0.2 using 0.1 M HNO 3 or 0.1 M NaOH. The samples were agitated on a Gerhard shaker at room temperature water bath at 120 rpm for 2 h. The solutions thus obtained were filtered through a 0.45 µm filter and analyzed for Cd 2+ concentration. Before metal analysis, the supernatants were acidified with HNO 3 . The residual adsorbent from the sorption experiment at selected conditions was dried to constant mass at 70 °C and characterized with advanced imaging and spectroscopic techniques.
Analytical methods. Total Cd 2+ concentration was measured with a flame atomic absorption spectrometer (FAAS) (Varian, AA28OFS) at 228.8 nm. The accuracy of the Cd 2+ analysis was validated with the reference material (CRM 142 R). The limit of Cd 2+ detection (LOD) was 0.07 mg/L, while the limit of Cd 2+ quantification (LOQ) was 0.21 mg/L. The functional groups on the surface of ALP-AGS and ALG, before and after Cd 2+ sorption, were analyzed in the range of 3800-400/cm using an FTIR spectroscope (Nicolet 6700, Thermo Scientific) equipped with a Smart Multi-Bounce HATR™. The surface morphology of adsorbents was examined with an LEO 1430VP scanning electron microscope (SEM) (Carl Zeiss). Qualitative and quantitative analysis of elemental surface composition (SEM-EDX) was performed with an energy-dispersive X-ray spectrometer (EDX, Quantax 200; detector: XFlash 4010, Bruker AXS, Berlin, Germany). Elemental mapping was performed on different micro-areas on the surface of ALP-AGS and ALG and the content of each element (in mass %) was averaged.
The Brunauer-Emmet-Teller (BET) specific surface area of adsorbents was determined by fitting the BET equation to the linear portion of the BET plot; the pore size distribution was calculated on the basis of the desorption plot of the N 2 adsorption-desorption isotherm using the Barret-Joyner-Halenda method (Micrometrics ASAP 2000, USA).
Calculations. The amount of Cd 2+ adsorbed onto the adsorbent in the equilibrium (q e , mg/g d.m.) was calculated according to the following formula: The efficiency (%) of Cd 2+ adsorption was calculated according to the following formula: where C 0 is the initial Cd 2+ concentration in the solution (mg/L), C e is the equilibrium Cd 2+ concentration (mg/L), m is the mass of adsorbent (g d.m.), and V is the volume of metal solution (L).
Equilibrium isotherm models. Three models were employed to interpret the adsorption isotherm data onto tested adsorbents: where: Freundlich model: K F is the constant in the Freundlich equation (L/mg), 1/n is a measure of adsorption intensity (-); Langmuir model: q max is the maximum monolayer adsorption capacity (mg/g d.m.), b is adsorption equilibrium constant (L/mg); Sips model: b is Sips isotherm constant related to the energy of adsorption (L/mg), n is a constant related to the grade of surface heterogeneity (-).
Kinetics model. The pseudo-second-order kinetic model was used to describe the kinetics of Cd 2+ adsorption onto adsorbents: where: q t is the amount of Cd 2+ adsorbed at a specific adsorption time (mg/g d.m.); k is a rate constant (g d.m./ (mg min)), and t is the adsorption time (min). The initial adsorption rate (r) was calculated as k‧q e 2 . The fitting of the adsorption and kinetics isotherms to experimental data and the adsorption (K F , q max , b, n) and kinetics parameters (q e , k) were calculated by the Levenberg-Marquardt optimization method included in the STATISTICA® software (version 13.3, TIBCO Software Inc.). The appropriateness of the adsorption and kinetic isotherm models was determined based on the sum of squared errors (SSE) and coefficient of determination (R 2 ).

Results and discussion
Optimization of ALP recovery from AGS. The yield of ALP-AGS was highest with an extraction temperature of 70 °C (Fig. 2a), an extraction time of 45 min (Fig. 2b), and 120 mL of 0.2 M Na 2 CO 3 /2.5 g MLSS (131 mg/g MLSS; significant differences, ANOVA, Tukey's HSD test, p = 0.05) (Fig. 2c). This time was shorter and the temperature of extraction was lower than in the original methodology presented by Lin et al. 17 , which creates some opportunities for decreasing the costs of ALP recovery from AGS in full-scale installations.
Seasonal and intra-cycle changes in ALP content in AGS. The concentration of ALP-AGS was investigated for over 3100 cycles of stable performance in two independently operated full-scale GSBRs at a municipal WWTP. In this period, no serious operational problems were reported, and wastewater composition did not vary significantly. Both GSBRs were operated at similar biomass concentrations, which caused the main operational parameters of the GSBRs, such as the organic loading rate, to be nearly identical. At the respective beginnings of spring 2018 and spring 2019, the ALP-AGS content reached nearly 100 mg/g MLSS and over 150 mg/g MLSS (Fig. 3a). These values were slightly higher than the ALP-AGS content reported in another study conducted in a full-scale WWTP 30 and similar to values observed in lab-scale reactors, which are usually higher than in fullscale systems 23 . For example, in lab-scale reactors fed with primary effluent from municipal wastewater, AGS contained 184 ± 18 mg VS ALP/g VSS 9 .
In this study, the ALP-AGS concentration in sludge increased significantly in the transition periods between winter and spring. During these periods, the temperature dropped to 9 °C and was the lowest of the entire experimental period (Fig. S1, Supplementary Materials). Bacteria may produce EPS as a strategy for survival in cold environments, as the presence of EPS significantly reduces cell lysis 31 . In pure strains of Lactobacillus paracasei, low temperatures increased the content of the high molecular weight fraction of EPS and the total amount of EPS produced 32 . Our results indicate that low temperatures also affect the ALP content in AGS-the highest yields of this biopolymer from waste sludge at the WWTP can be recovered at the end of winter. The average concentrations of ALP in biomass were higher in 2019 than in 2018; this may be explained by gradual granule maturation, which favors biopolymer production 33,34 . In a study by Huang et al. 35 , ALP were only found in mature granules.
Variations in the amounts and characteristics of ALP (mostly the molecular weight, MW) during the operational cycle of industrial-scale batch reactors have been reported 36 . In the present study, the amount of ALP varied considerably during the cycle, from about 50 to over 120 mg ALP/g MLSS, and two peaks of ALP content in biomass were observed, 1 and 2 h from the beginning of aeration (Fig. 3b).
ALP production has been reported to follow almost the same trend as bacterial growth 37 . Similarly, in the present study, the peak activity of bacteria, as indicated by the 16S rRNA levels in cells (results not shown), www.nature.com/scientificreports/ overlapped with the peaks of ALP production. During the introduction of wastewater to the reactors, the dissolved oxygen concentrations in the bulk liquid dropped significantly to below 0.5 mg/L in the GSBRs. Such low oxygen concentrations in the environment stimulate anoxic and anaerobic metabolism in the middle and core granule layers 38 . Therefore, the first peak in ALP formation may have resulted from anaerobic bacteria, such as phosphate-accumulating and glycogen-accumulating microorganisms, intensively producing ALP after the start of aeration 15 . As alginate contains glucuronic acid, it can act as a barrier to oxygen diffusion, limiting its transfer to the enzyme complexes in bacterial cells and protecting the metabolic activity of strict anaerobes 39 .
In the present study, ALP production started to increase in the 75th minute of the GSBR cycle and peaked at 2 h, which can be explained by an increase in ALP biosynthesis after the depletion of easily biodegradable acetate present in the wastewater introduced to the GSBR (data not shown). This explanation is consistent with a report of reduced ALP production in bacterial cells after the addition of sodium acetate to pure cultures of Azotobacter vinelandi cultivated on glucose-based media 40 . A similar pattern of ALP production has been reported in lab-scale reactors operated with an anaerobic/oxic/anoxic cycle and fed with wastewater containing sodium acetate 15 . In that study, ALP content was highest after 90 min of aeration. In the present study, the reduction in ALP levels in AGS at the end of the GSBR cycle can be explained by ALP lyase activity, which degrades ALP in the post-polymerization step 41 . This reduction in ALP levels is consistent with reports of batch tests, in which ALP production and its molecular weight dropped over time 42,43 . Practically speaking, our results indicate that, in a full-scale operation aimed at ALP recovery, the sludge must be discharged about 2 h after the start of aeration.
Adsorbent characteristics. ALP-AGS and ALG had a low surface area ( Table 1). The pore volume of ALP-AGS was about 24 times larger than that of ALG, but both adsorbents had similar pore diameters, the size of which indicated that they had microporous structures 44 . The surface area of ALG can vary considerably, depending on ALG conditions (e.g., the sodium alginate concentration, the type and concentration of gelation agent, etc.) or ALG modification. ALP-AGS that was produced with much more concentrated CaCl 2 (12.5%) than that in the present study had a much larger surface area (76.2 m 2 /g) and a slightly larger pore volume (0.0623 cm 3 /g), but the pores had a smaller diameter (0.0177 nm) 25 . Aziz et al. 45   . Similarly, Torres-Caban et al. 47 used SEM analysis to find that calcium alginate beads had a smooth surface without much porosity. At higher magnifications, the surface of the ALP-AGS was less regular and contained structures resembling crystalline outgrowths. An irregular and rough surface might indicate strong cross-links with Ca 2+34 and can be also attributed to water evaporation and surface shrinking 48 .

Adsorption of Cd 2+ onto ALG and ALP-AGS.
Cd 2+ adsorption onto the tested adsorbents was optimized in terms of adsorbent dosage, pH, initial Cd 2+ concentration, and sorption time.
Adsorbent dosage. At the lowest adsorbent dosage (0.7 g d.m./L), the process efficiency was higher for ALG (91%) than for ALP-AGS (75%) (Fig. 5a). This may be connected with the lower number of active sites on ALP-AGS. As the dosage of ALP-AGS was increased from 0.7 to 7.9 g d.m./L, the efficiency increased from 75 to 93%, which can be attributed to the presence of a greater number of adsorbent sites at a higher adsorbent dosage 50 . With adsorbent dosages in the range of 7.9-15.9 g d.m./L, the average Cd 2+ adsorption efficiency, at an initial metal concentration of 10 mg/L, was 94.3 ± 0.2% for ALG and 94.4 ± 0.9% for ALP-AGS. Based on the residual concentration of Cd 2+ in the solution after adsorption and the Cd 2+ removal efficiency, an ALP-AGS dosage of 7.9 g d.m./L was selected as optimum.
pH. The efficiency of Cd 2+ removal by ALG and ALP-AGS depended on the solution pH (Fig. 5b). The pH is known to be important for controlling metal adsorption, as it affects both the chemical properties of surface functional groups and the speciation of metal ions 51 . The best sorption effects with both sorbents were obtained with pH values in the range of 4-8; at these values, the average adsorption efficiency was 88.7 ± 0.9% for ALP-AGS and 97.0 ± 0.2% for ALG. Adsorption efficiency and adsorption capacity were lowest at pH 2-3. Kuczajowska-Zadrożna et al. 52 found that Cd 2+ adsorption onto ALG beads reached 91% at pH values ranging from 5.0 to 9.0, and that reducing the pH to 2.0 significantly decreased the adsorption efficiency to 23%. Mahmood et al. 53 reported that pH 6.0 was optimum for Cd 2+ adsorption onto ALG.
Changes in the pH can affect the charge of functional groups on the surface of the adsorbent, which, in turn, affects its capacity to adsorb metals. ALG sorbents consist mainly of guluronic and mannuronic acid units containing carboxyl and hydroxyl groups 54,55 . The pKa values of the carboxyl groups of the mannuronic and guluronic acid units in ALG are 3.38 and 3.65, respectively 54,55 . Although the individual constituents of the ALP-AGS were not analyzed in the present study, EPS that was recovered from AGS contained mainly proteins (≈ 87%), polysaccharides (≈10%), and humic acids (2.3%) 56 . Proteins are the main source of carboxyl, hydroxyl, and amine groups, while polysaccharides and humic acids are the sources of hydroxyl groups 56 . Although the content of humic acids in the EPS from AGS was lower than that of proteins, humic acids are high in carboxylic Table 1. Basic characteristics of the adsorbents (mean ± SD, n = 3). a Diameter of adsorbent beads after gelation. www.nature.com/scientificreports/ acids and phenols, implying that they might be used as chelating agents. Humic acids and proteins can form complexes with cationic metals that are beneficial for metal adsorption 57 . The negative surface charge of EPS at a pH range of 3 to 10 is related to the deprotonation of carboxyl (pKa ≈ 3.0), phosphoryl (pKa ≈ 6.5), amine (pKa ≈ 8.4), and hydroxyl (pKa ≈ 10.2) groups 56 . In the EPS from the granular sludge, carboxyl and hydroxyl groups were most abundant and were mainly responsible for Pb 2+ , Cd 2+ , and Zn 2+ adsorption 56 .  www.nature.com/scientificreports/ In the present study, all FTIR spectra (Fig. 6) contained absorption bands that indicated the presence of hydroxyl, ether, and carboxylic functional groups. From 3600 to 3000/cm, stretching vibrations of O-H bonds appeared, which are typical of polysaccharides 46 , and also N-H stretching vibrations of amino groups at ~ 3283/ cm in the ALG-AGS spectra, which could confirm the presence of proteins 27,58,59 . At 2952-2852/cm in ALP-AGS spectra, stretching vibrations of aliphatic C-H were observed. These bands can be assigned to fatty acids 60 , and their intensity was greater in the ALP-AGS spectra than in the ALG spectra (Fig. 6b). An additional band at ~ 3087/cm in the ALP-AGS spectra indicates the presence of aromatic structures (also the band at ~ 1540/cm). The latter band may also have a contribution from the vibrations of amide II groups, i.e., N-H bending (also visible at 1515/cm) and C-N stretching in proteins 60 . The presence of C-N bonds is, in turn, confirmed by the band at 1378/cm 58 . The presence of hydrophobic components does not affect the adsorbent's capacity for metal adsorption, but it can indirectly affect the location of polar hydrophilic groups responsible for metal adsorption 47 . The peaks at ~ 1730/cm are characteristic of C=O symmetric stretching in carboxylic acids. After Cd 2+ adsorption, this band is no longer visible in the ALG spectrum (Fig. 6a), which may indicate that Cd 2+ bonded with these acid groups. The bands at ~ 1590/cm (Fig. 6a) indicate the presence of ionic carboxylate salts.
From 1652 to 1630/cm in the ALP-AGS spectra, bands are visible that correspond to amide C=O and C-N stretching, N-H bending, C=C stretching in proteins, hydroxyl O-H stretching of polysaccharides 27,58,60 , and/ or COOgroups (Fig. 6b), which may indicate that these groups were more covalent in character. After Cd 2+ adsorption, the intensity of these bands in the ALP-AGS spectra decreased (Fig. 6b), which may indicate the participation of carboxylate groups in the formation of complexes with Cd 2+ . The bands at 1100-1000/cm correspond to the glycosidic bonds in the polysaccharide (C-O-C stretching) 56,61,62 .
The presence of bands at ~ 1410/cm may result from the overlapping of C-H group vibrations and COOgroup stretching 61 . This band can be assigned to the stretching vibration of C=O and the deformation vibration of -OH in carboxylate, alcohol, or phenol structures 56,63 , and its enhancement after adsorption is explained by the rearrangement of molecular bonds and therefore the formation of new bands. After Cd 2+ adsorption, both the ALG and ALP-AGS spectra have additional bands at ~ 1350/cm, and the bands at ~ 1410/cm have higher intensity, which may indicate both complexation of Cd 2+ and the presence of nitrates (the source of the Cd 2+ in the aqueous solution).
Thus, the high efficiency of Cd 2+ removal at pH values over 4.0 was related to metal complexation with negatively charged functional groups (especially COOH and OH groups) on the ALG and ALP-AGS surfaces according to these reactions: 2ALG-COO -+ Cd 2+ → (ALG-COO) 2 Cd 2+ and 2ALP-AGS-COO -+ Cd 2+ → (ALP-AGS-COO) 2 Cd 2+64 . The presence of monovalent species of Cd, which is due to the pH, can promote its complexation 65 . At pH < 6.0, it is present mainly as Cd 2+ . At pH > 6.0, the Cd 2+ content gradually decreases and other Cd-containing species appear, e.g., CdOH + , Cd 2 OH + 3 , and Cd(OH) 2 (s) 51,66 . Due to the amphiphilic character of EPS from AGS and the presence of abundant negatively-charged functional groups, metals can also be adsorbed via electrostatic attraction, ion exchange, or surface precipitation 56,67 . Contact time and adsorption kinetics. The effect of contact time on Cd 2+ adsorption onto ALG and ALP-AGS is shown in Fig. 7. The amount of Cd 2+ adsorbed at a specific adsorption time (q t ) indicates that, as the contact time was increased from 5 to 60 min, Cd 2+ adsorption increased, and then between 60 and 180 min, the adsorption curve was flat. Similarly, Liu et al. 56 observed very quick metal (Pb 2+ , Cd 2+ , Zn 2+ ) adsorption onto EPS recovered from AGS cultivated in a lab-scale sequencing batch reactor. The metal adsorption sharply increased within the first 10 min, and progressively slowed until saturation after about 60 min. The fast adsorption onto EPS could be related to metal interactions with functional groups of proteins. The R 2 and SSE values indicate that Cd 2+ adsorption onto both types of adsorbents at different contact times was well described by a pseudo-second-order kinetics model (Fig. 7) 54 .
The good fit of the model to the data indicates that the rate-limiting step in Cd 2+ adsorption is chemisorption, which in the case of ALG and ALP-AGS, was due to complexation and electrostatic attraction between Cd 2+ and negatively charged groups, as well as ion exchange between Cd 2+ and other cations on the adsorbents' www.nature.com/scientificreports/ surfaces, e.g., Ca 2+ . During the initial stage of adsorption, Cd 2+ removal was rapid, as shown by the initial rate of adsorption (r), which was higher for ALP-AGS than for ALG. As time elapsed, the adsorption process slowed, due to the saturation of the available surface-active sites. For both adsorbents, equilibrium conditions for Cd 2+ removal were reached within 60 min (Fig. 7).
Initial Cd 2+ concentration and adsorption isotherms. The effects of the initial Cd 2+ concentration (5-200 mg/L) on the metal removal efficiency and the adsorption capacity of the ALG and ALP-AGS are shown in Fig. 8a-d respectively.
The initial Cd 2+ concentration affected the efficiency of adsorption with ALG and ALP-AGS. At initial Cd 2+ concentrations in the range of 5-15 mg/L, the average Cd 2+ removal efficiency slightly increased from 96.7 to 98.9% for ALG and from 92.1 to 94.6% for ALP-AGS. However, at initial concentrations above 15 mg/L, the efficiency gradually decreased to 73.5% (ALG) and 79.9% (ALP-AGS) (Fig. 8c, d). This means that the presence of available adsorption sites onto the adsorbents was high up to the relevant Cd 2+ concentration.
The relationship between the amount of Cd 2+ adsorbed by ALG and ALP-AGS and the concentration of Cd 2+ remaining in the solution was evaluated with the Freundlich, Langmuir, and Sips models. Of these three adsorption models, the Sips model had the highest R 2 values and the lowest SSE values ( Table 2), indicating that it was the most suitable for describing Cd 2+ adsorption onto tested adsorbents. The q max is one of the principal criteria for the suitability of a particular adsorbent for metal removal. According to the Sips model, the q max values were 29.3 mg/g d.m. for ALG and 29.5 mg/g d.m. for ALP-AGS. With comparable q max , a lower value for energy of adsorption (b) was for ALP-AGS. With the Langmuir model, the q max value for ALP-AGS was similar to this calculated from the Sips model. Thus, ALP-AGS has the potential to serve as a substitute for commercial ALG. These q max values are similar to those reported from other studies of Cd 2+ adsorption with ALG adsorbents, e.g., ALG-calcium carbonate beads: q max, Langmuir = 10.2 mg/g 53 , q max, Langmuir = 37.8-68.9 mg/g 68 , q max, Langmuir = 38.05 mg/g and q max, Sips = 38.65 mg/g 47 . Liu et al. 56 found that EPS extracted from AGS exhibited a high adsorption capacity for Cd 2+ (1470 mg/g) that was much higher than for ALP-AGS and conventional biosorbents. The differences in adsorption capacities can be related to the methods used for the biopolymers' recovery from the sludge matrix and their purification, which can provide more efficient exposure to the negatively charged sites. In the present study, ALP recovered from AGS was not purified.  69 . Based on the values of heterogeneity index (n), adsorption of Cd 2+ by ALP-AGS seems to occur onto more uniform and homogenous active sites compared to ALG 69 . A degree of homogeneity/heterogeneity of the adsorption sites can depend on the number of functional groups with the same capacity for adsorption. Cd 2+ with low bonding strength is mainly adsorbed by carboxyl groups in the ALG molecule. Due to the conformational changes in the saccharide chains of G and M blocks, all carboxyl groups are not easily available, which affects the adsorption process 52 . Within the egg-box structure of ALG, the G-block carboxyl groups are less readily available to the metal ions, whereas the M-block carboxyl groups can more easily interact with Cd 2+70 . In contrast to ALG, active sites on ALP-AGS, come not only from polysaccharides but also from proteins or humic acids 71 . The investigations of metal adsorption onto individual EPS components (proteins, humic acids, and polysaccharides) recovered from different sludges have demonstrated that proteins exhibited the highest adsorption capacity for Cd 2+ removal, while the polysaccharides were less efficient 71 . Thus, it might suggest that some irregularities in active sites onto polysaccharides in ALP-AGS can be compensated by active sites from other polymers co-existing in ALP.
Chemical composition of alginate adsorbents after Cd 2+ adsorption. SEM-EDX analysis of adsorbents after Cd 2+ adsorption revealed that their composition had undergone some changes during the adsorption (Figs. 8, S2) process. Sorption of Cd 2+ caused a decrease in the share of most elements analyzed with SEM-EDX, especially C (which may be related to changes in the carbon structure of alginate), as well as Ca (which may be related to the ion exchange mechanism of Cd 2+ adsorption) 72 . Alkali metals play the role of ion exchange in the process of metal adsorption. Among them, Ca ions play an important role in the ion exchange of metal cations under medium acidic conditions 73,74 . In the present study, the experiment on Cd adsorption was performed at pH 5.0, and Ca release was observed from both ALG and ALP-AGS. Similarly, Ablough et al. 75 found that after the adsorption of Pb on hybrid beads of chitosan and sodium alginate, the Pb peak appeared in the EDX spectra, while the Ca peak disappeared, indicating that Ca was completely removed from the adsorbent and ion exchange was the main mechanism of Pb adsorption. Bée et al. 76 related the binding of Pb to Ca-alginate beads to the release of Ca and found that the amount of Ca remaining in the biopolymer bead was very small when the sorbent was saturated with Pb. Some divalent heavy metal ions with high reactivity can exchange Ca ions while maintaining the structure of the crosslinked adsorbent 55 without unraveling the polymer.
The surface EDX mapping of the ALG and ALP-AGS proved the realization of the Cd 2+ adsorption process, with comparable mass % of Cd 2+ onto ALP-AGS (3.92 ± 1.25 mass %) and ALG surfaces (3.20 ± 0.43 mass %). These results correspond well with the maximum adsorption capacities estimated from Langmuir and Sips models ( Table 2). The images obtained through mapping elements in SEM-EDX for Cd 2+ exhibit uniform metal distribution on the surface of both adsorbents (Fig. 9).

Conclusions
This study showed that AGS from a full-scale WWTP is a rich source of ALP, which can be used for the effective adsorption of Cd 2+ . ALP content in the AGS from the full-scale facility increased in the transition period between winter and spring, reaching over 150 mg/g MLSS. In the batch reactor cycle, ALP content was highest 1 h after the start of aeration, about 2 times higher than at the end of the cycle. ALP-AGS has a low total pore volume and surface area, and its sorption properties were determined by the presence of carboxylic and hydroxyl groups. The Cd 2+ removal mechanism was governed by chemisorption and a monolayer of Cd 2+ on ALP-AGS surface was created. The most efficient Cd 2+ removal was observed at an adsorbent dosage of 7.9 g d.m./L. ALP-AGS, pH ranging from 4 to 8, and 60 min. Under these conditions, the mass percentage of Cd 2+ adsorbed onto ALP-AGS was 3.92%, while that adsorbed onto commercial ALG was 3.20%. ALP-AGS can serve as an attractive substitute www.nature.com/scientificreports/ for commercial ALG in the elimination of toxic Cd 2+ from the environment, and sustainable ALP recovery from waste AGS can be implemented as part of comprehensive strategies for utilizing wastewater in a circular economy.

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